J Appl Phycol DOI 10.1007/s10811-017-1280-6
Evaluation of domestic wastewater treatment using microalgal-bacterial processes: effect of CO2 addition on pathogen removal Graziele Ruas 1 & Mayara Leite Serejo 1 & Paula Loureiro Paulo 1 & Marc Árpád Boncz 1
Received: 4 May 2017 / Revised and accepted: 14 September 2017 # Springer Science+Business Media B.V. 2017
Abstract Microalgal-bacterial processes represent a sustainable and cost-effective biotechnology able to promote efficient wastewater treatment, including natural pathogen removal (disinfection), as well as being able to perform CO2 uptake and biogas upgrading. In this context, the influence of CO2 supply from a synthetic gas mixture (30% v/v CO2) on the removal of pathogens (Pseudomonas, enterococci, and Escherichia coli) and total coliforms during secondary domestic wastewater treatment by a microalgal-bacterial symbiosis in a 180-L high-rate algal pond (HRAP) was investigated. The supply of CO 2 in the HRAP positively influenced the Pseudomonas aeruginosa removal, with the removal efficiency increasing from 97.4% (1.6 log) to 99.6% (2.5 log) without and with CO2 supply, respectively. Likewise, the total coliform removal efficiency rose from 88.7% (1.1 log) to 99.4% (2.8 log). On the other hand, the effect of CO2 supply on enterococci (99.7% and 2.6 log) and Escherichia coli (98.6% and 2.2 log) removal was negligible.
Keywords Biofuel . Biogas upgrading . Chlorella vulgaris . Disinfection . Wastewater treatment
Electronic supplementary material The online version of this article (https://doi.org/10.1007/s10811-017-1280-6) contains supplementary material, which is available to authorized users. * Mayara Leite Serejo
[email protected]
1
Faculty of Engineering, Architecture and Urbanism and Geography, Federal University of Mato Grosso do Sul, Avenida Costa e Silva s/n, Campo Grande, Brazil
Introduction The inappropriate disposal of domestic and industrial wastewaters into the environment causes many pollution problems such as eutrophication or oxygen depletion in water bodies, besides public health problems via ingestion or contact with water contaminated by pathogenic bacteria (Abdel-Raouf et al. 2012; Cai et al. 2013; Posadas et al. 2015). According to the World Health Organization, approximately 1.8 billion people worldwide drink water contaminated with fecal matter, especially in underdeveloped countries, due to a lack of adequate sanitation such as water and wastewater treatment (WHO and UNICEF 2014; WHO 2015). In this context, diseases such as diarrhea and severe bacterial infections can be spread by pathogen microorganisms such as Escherichia coli, Pseudomonas sp., and Enterococcus sp. (Bitton 2005; Hocquet et al. 2016). Furthermore, great concern is currently focused on the resistance of these pathogens to antibiotics (normally called multidrug-resistant pathogens or antibioticresistant bacteria), as approximately 80% of the bacteria in wastewater have been shown to have resistance to some antibiotic (Bouki et al. 2013; Rizzo et al. 2013; Zabawa et al. 2016). Contextually, E. coli, Pseudomonas aeruginosa, and enterococci (especially Enterococcus faecalis and Enterococcus faecium) are the major multidrug-resistant bacteria identified in hospital wastewater (Hocquet et al. 2016). Conventional biological wastewater treatment technologies unfortunately present some techno-economic limitations such as high aeration costs (activated sludge reactors) and poor pathogen and nutrient removal (anaerobic reactors) (de la Noüe et al. 1992; Posadas et al. 2014a). Furthermore, despite the higher removal efficiency of pathogenic bacteria (up to 5.7 log for E. coli), stabilization ponds for tertiary treatment require large areas to perform the treatment (Picot et al. 1992; Dias et al. 2014). Microalgal-bacterial symbiosis processes represent a sustainable
J Appl Phycol
and cost-effectivebiotechnology thatis abletoremove pathogens and nutrients at a lower energy cost and a smaller area requirement (Picot et al. 1992; Muñoz and Guieysse 2006). This biotechnology is characterized by the photosynthetic conversion of CO2 to microalgal biomass concomitantly with assimilation of nutrients by cells producing O2, which is used for the oxidation of organic pollutants to CO2 and also for nitrification reactions (Muñoz and Guieysse 2006). The microalgal-bacterial process usually occurs in high-rate algal ponds (HRAPs), first implemented in the mid-1950s in California for domestic wastewater treatment (Oswald 1988). Under optimal conditions, HRAPs can treat higher concentrations of organic matter than waste stabilization ponds (about three to five times), at the same time reducing the surface needed by a factor of 5 (Picot et al. 1992; Racault and Boutin 2005; Muñoz and Guieysse 2006). Furthermore, natural disinfection can occur in HRAPs due to different factors such as solar radiation, oxygen supersaturation, alkaline pH, temperature, and bactericidal effects produced by microalgae, among others (Davies-Colley et al. 1997; Craggs et al. 2004; Muñoz and Guieysse 2006; Reinoso et al. 2008; Ouali et al. 2014). Finally, the biomass produced in HRAPs can be used as a biofertilizer and/or as a material for biofuel production. Typical raw domestic wastewater is normally carbonlimited for microalgal growth (Craggs et al. 2012; Posadas et al. 2013; Sutherland et al. 2015). In this context, the CO2 supply in HRAPs can prevent carbon limitation and pH inhibition of microalgal growth, and can avoid ammonia toxicity and stripping as well as phosphorus precipitation in the mixed liquor (Heubeck et al. 2007; Park and Craggs 2010; Craggs et al. 2012; Posadas et al. 2015). Furthermore, upgrading of biogas to produce biomethane can be performed in HRAPs via CO2 and H2S removal by microalgae and sulfur-oxidizing bacteria, respectively (Toledo-Cervantes et al. 2013; Serejo et al. 2015). On the other hand, Garcia-Gonzalez et al. (2007) reported the inhibition of bacterial activity in aqueous medium with CO2 addition (CO2 bactericidal action). Thus, the use of CO2 (from gas or biogas) can be an advantageous alternative for disinfection in HRAPs. The number of studies focused on the concomitant removal of P. aeruginosa, enterococci, E. coli, and total coliforms by microalgal-bacterial processes in HRAPs treating raw domestic wastewater (RDW), with or without CO2 addition, is unfortunately few. In one study, Bahlaoui et al. (1997) reached a removal efficiencyof≈91.6%(≈1.1log)ofP.aeruginosafromrawsewage in a 48-m2 HRAP, operating with a hydraulic retention time (HRT) of 8 days, without CO2 supply. In another study, also without the addition of CO2, Awuah et al. (2002) achieved an enterococci removal efficiency from anaerobically pre-treated domestic wastewater of ≈ 2.0 log in an algae-based treatment pond with 7day HRT. More recently, Heubeck et al. (2007) and Posadas et al. (2015)recordedanE.coliremovalefficiencyofapproximately2.0 and 1.0 log, both in HRAPs with CO2 supply, treating raw and anaerobically digested sewage, respectively.
The main objective of this work was to study the influence of the addition of CO2 on the removal of pathogenic bacteria (P. aeruginosa, enterococci, and E. coli) and total coliforms from RDW by using a microalgal-bacterial symbiosis process in a 180-L HRAP. Furthermore, the potential carbon and nutrient removal from the wastewater, the biomass settleability, and the dynamics of the microalgae population in the HRAP were also investigated.
Materials and methods Experimental setup and operational conditions The experimental setup consisted of a 180-L pilot HRAP with an illuminated surface of ≈ 1.2 m2 (110 cm long × 110 cm wide) and 15-cm culture depth. Four submerged pumps with a nominal flow rate of 540 L h−1 (Sarlo Better B500, Brazil) were located at the bottom of the reactor (one in each corner of the reactor) in order to promote complete mixing of the culture, which resulted in a liquid recirculation velocity of 20 ± 2 cm s−1 (Fig. 1), comprising the hydraulics characteristics of conventional HRAP setups (Oswald 1988; Andersen 2005; Muñoz and Guieysse 2006; Craggs et al. 2014). Almost no dead zones were observed in the HRAP. An acrylic 2.0-L (Ø = 5.0 cm; length = 100 cm) absorption column (AC) was introduced inside the HRAP, horizontally inclined at ≈ 15°. The bottom of the AC was opened to allow entry of liquid and gas, while the top had two separate connections for liquid and gas output. Liquid recirculation in the AC was carried out with a pump (Watson Marlon 505U, UK) and the liquid returned back to the HRAP. After leaving the HRAP, the mixed liquor passed through an 8-L settler operating at a hydraulic retention time (HRT) of 12.0 ± 1.0 h. The experiment was operated in a greenhouse (solar radiation) located at the University of Mato Grosso do Sul (Brazil) for 51 days at ≈ 36 °C. The spectral signature of the greenhouse in terms of absorbance and transmittance is presented in Fig. S1 (Supplementary Material). Gas and raw domestic wastewater A synthetic gas mixture (SGM) composed of CO2 (30%) and N2 (70%) (White Martins, Brazil) was used as a source of CO2 for the pathogen removal efficiency experiments in the HRAP–AC system. RDW was collected twice per week from the WWTP and stored in a 200-L feed tank with a submerged pump with a nominal flow rate of 500 L h−1 (Guangdon Minjiang NS 802, China) for agitation in order to prevent sedimentation of suspended solids. The HRAP was initially filled with 166 L raw domestic wastewater and fed with RDW in continuous mode using a pump (Watson Marlon 505U, UK) applying a HRT of 5.0 ± 0.5 days, which was chosen
J Appl Phycol
Fig. 1 Schematic diagram of the experimental setup for continuous photosynthetic pathogen, total coliform, and nutrient removal
based on typical HRTs reported for domestic wastewater treatment in HRAPs (Arbib et al. 2013; Posadas et al. 2015).
Microorganisms and culture conditions The HRAP was inoculated with 12 L of a 0.9-g total suspended solids (TSS) L−1 culture of microalgae (≈ 97% Chlorella vulgaris) and 2 L of an 8.1-g TSS L−1 nitrifying– denitrifying aerobic activated sludge from Mato Grosso do Sul domestic wastewater treatment plant (WWTP). The initial analyzed TSS concentration in the cultivation broth of the HRAP was 0.15 g L−1. Microalgae were collected and previously acclimated to the RDW prior to HRAP inoculation according to Serejo et al. (2015).
Continuous experiments The experiment was performed considering two operational periods in order to study the influence of CO2 dosing on the removal efficiency of pathogens (Pseudomonas aeruginosa, enterococci, and Escherichia coli) and total coliforms, using the microalgalbacterial symbiosis process. Stage I was carried out without SGM addition(Table1).Ontheotherhand,atstageII,SGMwassparged continuously into the AC from the bottom, at a flow rate of 5 mL min−1, while the liquid recirculation rate was 50 mL min−1 with the purpose of maintaining a liquid recirculation/gas (L/G) ratio of 10 (Serejo et al. 2015). Each operational stage was maintained for approximately 25 days (≈ 5 times the HRT of the HRAP). The average initial concentrations of soluble chemical oxygen demand (COD), total organic carbon (TOC), inorganic carbon (IC), total nitrogen (TN), N–NH4+, N–NO2−, N–NO3−, phosphorus (P), P. aeruginosa, enterococci, E. coli, total coliforms, pH, and TSS of the RDWare summarized in Table 2.
Liquid samples of 500 mL were drawn twice a week from the influent (HRAP input, after the peristaltic pump) and effluent (settler output, Fig. 1) to monitor the concentration of soluble (filtration through 0.20-μm nylon filters prior to analysis) COD, TOC, IC, TN, N–NH4+, NO2−, NO3− P, and non-soluble pathogens and TSS. The samples were filtered for the purpose of simulating biomass harvesting from the HRAP with a membrane module (Posadas et al. 2015). Likewise, liquid samples of 200 mL were drawn from the cultivation broth twice a week to monitor the TSS concentration. The ambient and cultivation broth temperatures, light intensity, pH, and dissolved oxygen (DO) concentration were measured daily. Furthermore, the daily evaporation rate was determined from the difference between the wastewater influent and effluent flow rates (Serejo et al. 2015). For this, the influent flow rates were measured by calibrating the pump three times a week, while effluent flow rates were obtained daily by collecting the effluent in a graduated cylinder. Biomass harvesting was performed twice a week, while the CO2 output concentration was measured daily. The input and output flow rates of the gas were also measured to accurately determine CO2 removal. Finally, sampling was always conducted at 10:00 a.m. throughout the entire experimental period. The calculation of soluble TOC, IC, total inorganic carbon, COD, TN, N–NH4+, and P removal efficiencies; the surface organic loading rate (OLR) applied in the HRAP; the total suspended solid removal efficiency of the settler (TSSSet–RE); biomass productivity; and C, N, and P mass balance are described in the Supplementary Material section 2 (S2).
Analytical procedures The TOC and IC concentrations were determined using a total organic carbon analyzer TOC-VCSH (Shimadzu, Japan). Total nitrogen concentration was measured using Hach kits, method
J Appl Phycol Table 1 Main operational parameters during the two different operational stages
Conditions
Unit
Stage I
CO2 addition Sampling month Elapsed time
–
No
Yes
– day
January 25
February 26
HRT Influent flow rate
day L day−1
5.0 ± 0.4 36.4 ± 1.6
4.9 ± 0.6 37.7 ± 2.6
Light intensity
W m−2
82.2 ± 19.8 231.6 ± 50.5 13.4 ± 0.2
66.4 ± 19.8 200.4 ± 51.5 12.9 ± 0.3
36.1 ± 3.4
35.0 ± 4.5
−2
Average hourly irradiation intensity Light time Air temperature
Wm h day−1 °C
Stage II
Data shown are mean ± standard deviation, n = 8
10071, with a DR 6000 spectrophotometer (Hach, Canada). N–NH4+ and pH were measured using ammonia and pH electrodes, respectively, in combination with a Orion Five Star multiparameter meter (Thermo Scientific, The Netherlands). N-NO3−, N-NO2−, and P-PO43− were analyzed by means of a Dionex UltiMate ICS 1100 with a IonPac AG19/AS19 column (Thermo Scientific, USA). Pseudomonas aeruginosa, enterococci (Enterococcus faecium plus Enterococcus faecalis), E. coli, and total coliforms were determined using the methodology and Pseudalert®, Enterolert®, and Colilert® kits, respectively (IDEXX Laboratories, USA). All analyses, including COD and TSS, were carried out according to Standard Methods (APHA 2012). DO and temperature in the HRAP were measured using a 9500 DO2 Meter oximeter (Jenway, UK). The light intensity was recorded with a LX-101 lux meter (Lutron Corporation, Taiwan), and average hourly irradiation intensity, recorded at a weather station located in Campo Grande city (PCD 31950; 20° 30′ 07.2″ S and 54° 37′ Table 2 Initial physical-chemical and biological characterization during the two different operational stages
08.4″ W), 1 km away from the HRAP, were obtained from the Brazilian National Space Research Institute (INPE 2017). The spectral signature of the greenhouse was determined using a DR6000 UV VIS spectrophotometer (Hach, USA). The identification of microalgae was carried out by microscopic examination (Olympus BX41, USA) of microalgae samples (fixed with Lugol’s solution at 5% and stored at 4 °C prior to analysis) according to Sournia (1978). The determination of the C and N content of the microalgal-bacterial biomass was performed using a vario EL cube (Elementar, Germany). The CO2 output measurement was performed by a portable 0– 100% CO2 analyzer (Gas Expert, Brazil). Statistical treatment The results were evaluated using an analysis of variance (ANOVA) with a Fisher’s least significant difference (LSD) test using a 95% confidence level. A Pearson correlation
Parameters
Unit
Stage I
Stage II
COD OLR TOC IC TIC TN NH4+ NO2− NO3− P pH Pseudomonas aeruginosa Enterococci
mg COD L−1 g COD m-2 day−1 mg C L−1 mg C L−1 mg C L−1 mg N L−1 mg N-NH4+ L−1 mg N-NO2− L−1 mg N-NO3− L−1 mg P-PO43- L−1 – MPN (100 mL)−1 MPN (100 mL)−1
134 ± 47 4.3 ± 1.6 52 ± 19 117 ± 13 117 ± 13 32 ± 8 31 ± 5 5.7 ± 4.8 1.1 ± 0.4 3.2 ± 0.3 7.8 ± 0.3 (1.4 ± 0.2) × 106 (3.5 ± 0.9) × 104
490 ± 112 16.2 ± 3.7 185 ± 44 166 ± 35 206 ± 35 43 ± 7 43 ± 3 5.7 ± 4.1 2.0 ± 0.4 3.8 ± 0.5 7.7 ± 0.3 (1.0 ± 0.5) × 106 (1.8 ± 2.0) × 104
Escherichia coli Total coliforms
MPN (100 mL)−1 MPN (100 mL)−1
(9.6 ± 9.7) × 103 (1.6 ± 1.5) × 107
(5.7 ± 4.6) × 103 (3.9 ± 4.8) × 107
The average value (n = 4) is shown for all parameters. Variations are standard deviation MPN most probable number
J Appl Phycol
analysis (n = 4) was performed to verify the effect of CO2, DO, sunlight intensity, and pH on the pathogen and total coliform removal efficiencies.
Table 3 pH, DO concentration, temperature, and evaporation rate of the cultivation broth during the two different operational stages Parameters
Unit
Stage I
Stage II
pH
–
DO
mg O2 L−1
7.7 ± 0.2 a 7.3–8.0 b 8.5 ± 1.2 a 7.7–9.4 b
6.8 ± 0.3 a 6.3–7.5 b 6.3 ± 1.6 a 4.1–7.7 b
Temperature
°C
Evaporation rate
L m−2 day−1
29.1 ± 2.6 a 25.2–32.9 b 3.0 ± 0.6 a 1.5–4.0 b
28.0 ± 1.4 a 26.0–31.3 b 2.8 ± 0.9 a 1.6–3.9 b
Results Operational and environmental conditions in the HRAP Similar operational conditions such as HRT, influent flow rate, light intensity, average hourly irradiation intensity, light time, and ambient temperature were recorded in stage I (no synthetic gas supply) and II (with synthetic gas supply) during the 51 days of HRAP operation (p < 0.05) (Table 1). Although the average light intensity and average hourly irradiation intensity both seemed to decrease from stages I to II, the statistical analysis showed similarity between the two stages (p = 0.09), considering a confidence level of 95% (p < 0.05). Seasonal variation in the RDW resulted in increasing influent COD, TOC, IC, TN, and N-NH4+ concentrations during stage II (Table 2). However, a comparable RDW C/N ratio was obtained for both stages I (6.1 ± 0.5) and II (6.4 ± 0.8). During stage I, the pH in the cultivation broth remained relatively low at 7.7 ± 0.2 (Table 3), while the gas mixture supply at stage II resulted in a decrease of pH to 6.8 ± 0.3. Similarly, the DO concentration decreased from 8.5 ± 1.2 to 6.3 ± 1.6 mg O2 L−1 in stages I and II, respectively. On the other hand, comparable cultivation broth temperatures and evaporation rates of 29.1 ± 2.6 and 28.0 ± 1.4 °C and 3.0 ± 0.6 and 2.8 ± 0.9 L m−2 day−1 were found for stages I and II, respectively. Organic matter, carbon, and nutrient removal efficiencies Although the OLR increased from 4.3 ± 1.6 to 16.2 ± 3.7 g COD m−2 day−1 (Table 2), both COD and TOC–REs increased from 67 ± 11% (stage I) to 88 ± 9% (stage II) and from 59 ± 13 to 80 ± 13%, respectively. Similar IC–REs were found during stages I (70 ± 9%) and II (72 ± 4%), while the TIC–REs increased from 70 ± 9 to 78 ± 3% during stage II (Table 4). The carbon mass balance revealed that assimilation into biomass was the main C removal mechanism during stage I, whereas unfortunately, CO2 stripping was the principal mechanism during stage II. In this context, merely 9 ± 1% of CO2 removal efficiency was achieved in the AC. Assimilation into biomass was the only mechanism of N removal in the HRAP in both stages, where high nitrification activity was recorded regardless of the studied stage (≈ 97% of N–NH4+–REs). TN–REs decreased from stage I (39 ± 6%) to stage II (27 ± 11%), but this reduction was correlated with the increasing RDW nitrogen influent concentration. During stages I and II, the effluent N–NO 2 − (11.9 ± 4.7 and
The average value (n = 8) is shown for all parameters a
Average ± standard deviation
b
The range minimum–maximum measured value of each parameter
21.3 ± 0.8 mg N–NO2− L−1, respectively) and N–NO3− (10.7 ± 5.9 and 5.5 ± 0.6 mg N–NO3− L−1) concentrations corresponded to about all the effluent TN recorded. Assimilation into biomass was also the main phosphorus removal mechanism in both stages, as the low pH prevented P precipitation. Thus, P–REs of only 16 ± 4 and 8 ± 1% were recorded during stages I and II, respectively. Influence of CO2 source on removal efficiency of pathogens and total coliforms Similar REs were found for both enterococci and E. coli during both stages (99.8 ± 0.0%; 2.7 ± 0.1 log–REs during stage I and 99.7 ± 0.2%; 2.6 ± 0.3 log–REs during stage II for enterococci, and ≈ 98.6%; ≈ 2.2 log–REs for E. coli during both stages). Also, in stage I, higher total coliform REs were recorded (88.7 ± 1.5%; 1.1 ± 0.1 log–REs) than in stage II (99.4 ± 0.6%; 2.8 ± 1.2 log–REs). Pearson’s r correlations of 0.934, 0.484, 0.610, and 0.160 were obtained for CO2 addition (p < 0.05), DO (p < 0.05), sunlight intensity (p > 0.05), and pH (p > 0.05). However, a P. aeruginosa RE of 97.4 ± 0.5% (1.6 ± 0.1 log–REs) was found during stage I (Fig. 2a, b), while during stage II, with CO2 addition, the P. aeruginosa RE increased to 99.6 ± 0.4% (2.5 ± 0.5 log–REs). Pearson’s r correlations of 0.961, 0.439, 0.699, and 0.357 were obtained for CO2 addition (p < 0.05), DO (p < 0.05), sunlight intensity (p > 0.05), and pH (p > 0.05). Productivity and settleability of biomass, microalgae population, and elemental composition The TSS concentrations in stage I (0.12 ± 0.05 g L−1) remained comparable to those recorded at stage II (0.11 ± 0.04 g L−1). Likewise, similar biomass productivities were found during stages I (4.2 ± 1.1 g m−2 day−1) and II
J Appl Phycol Table 4 COD, TOC, IC, TIC, TN, N-NH4+, and TSSSet removal efficiency; TSS concentration; and biomass productivity during the two different operational stages TSS concentration (g L−1) Biomass productivity (g m−2 day−1)
Stage Removal efficiency (%) COD
TOC
IC
TIC
N-NH4+ P
TN
I
67 ± 11 59 ± 13 70 ± 9 70 ± 9 39 ± 6
II
88 ± 9
96 ± 3
80 ± 13 72 ± 4 78 ± 3 27 ± 11 97 ± 3
TSSSet
16 ± 4 71 ± 2 0.12 ± 0.05
4.2 ± 1.1
8±1
4.1 ± 1.3
96 ± 2 0.11 ± 0.04
The average value (n = 4) is shown for all parameters. Variations are standard deviation
(4.1 ± 1.3 g m−2 day−1), and similar algal-bacterial C and N contents (on a dry weight basis) of 46.8 and 8.3% were found at inoculation, 45.6 and 8.8% at stage I, and 46.6 and 8.6% at stage II, respectively. Cell populations changed drastically from stage I to stage II though. At stage I, about 91% of C. vulgaris and 6% of Microspora sp. were recorded (Fig. 3), but the percentage of cell numbers changed drastically, to 46 and 52% of C. vulgaris and Microspora sp., respectively, at stage II. At the same time, the TSSSet–REs increased from 71 ± 2 (stage I) to 96 ± 2% (stage II).
Discussion Operational and environmental conditions in the HRAP The RDW C/N ratio obtained for stages I and II was much higher than the typical RDW C/N ratios of ≈ 3.0 (Posadas et al. 2013; Sutherland et al. 2015), but consistent with the optimum C/N ratio (≈ 5.6) reported for microalgae growth (Borowitzka and Borowitzka 1988). The relatively low pH of the cultivation broth during stages I and II was due to the high RDW buffer capacity, as a consequence of the high IC concentrations (117–166 mg C L−1) (Posadas et al. 2015). Contextually, this is advantageous since acidic environments (pH 5–7) are preferred by eukaryotic microalgae (such as C. vulgaris and Microspora sp.), while alkaline environments (pH 7–9) are preferred by cyanobacteria (Razzak et al. 2013). On the other hand, the decreasing DO concentration from stages I to II suggests higher bacterial oxidation activity influenced by the increasing
B
100
Organic matter, carbon, and nutrient removal efficiencies Approximately 14 g COD m−2 day−1 was removed in the HRAP during stage II, higher than the COD–RE recorded by Matamoros et al. (2015) (≈ 66%; ≈ 8 g COD m−2 day−1) in a 470-L HRAP treating domestic wastewater at a lower influent OLR (≈ 12 g COD m2 day−1) and temperature (≈ 28 °C), but a higher pH (> 8). Furthermore, the effluent COD concentrations obtained during stages I (45.1 ± 23.0 mg COD L−1) and II (52.2 ± 22.1 mg COD L−1) remained lower than the maximum COD concentration acceptable for wastewater discharge into the environment of 125 mg COD L−1, according to Directive 91/271/ECC (European Union 1991). On the other hand, the poor removal efficiency of CO2 in the absorption column was due to the low CO2 solubility at the 4.0 3.0
90 (log)
Removal Efficiency (%)
A
RDW influent organic matter concentration; however, DO concentrations always corresponded to aerobic conditions (above 2.0 mg O2 L−1) (Posadas et al. 2014a). The pH and DO range obtained in our experiments are in agreement with the results found by Heubeck et al. (2007), Park and Craggs (2010), and Posadas et al. (2014a, b) of 6.3–9.2 and 0.5– 7.3 mg O2 L−1 for pH and DO, respectively, when treating domestic wastewater in HRAPs. Furthermore, the temperatures during the two experimental stages were within the optimum temperature range for many green microalgae of between 28 and 35 °C (Soeder et al. 1985); however, recorded evaporation rates were higher than those estimated for HRAPs under outdoor conditions in tropical climates (≈ 1.3 L m−2 day−1), as a result of high turbulence in the HRAP (Guieysse et al. 2013).
80
Pseudomonas Enterococcus
2.0
E. coli 1.0
70
Total coliform
0.0
60 I
II Stage
I
II Stage
Fig. 2 Removal efficiency in percentage (%) (a) and logarithmic (log) (b) of Pseudomonas aeruginosa, enterococci, Escherichia coli, and total coliforms during the two different operational stages. The average value (n = 4) is shown for all parameters. Variations are standard deviation
J Appl Phycol Chlorella vulgaris
Microspora sp.
Other species (< 3%)
Microalgae Assemblage
100%
80%
60%
40%
20%
0% Inoculation
Stage I
Stage II
Fig. 3 Dynamics of microalgae population in percentage of number of cells during the inoculation and the two experimental stages evaluated
low pH recorded (Serejo et al. 2015). Contextually, a higher CO2–RE of ≈ 50% was reported by Kao et al. (2012) during biogas upgrading by Chlorella sp. MB-9 in a 50-L outdoor photobioreactor using a basic synthetic medium. A relatively high TN–RE of ≈ 59% (corresponding to elimination of ≈ 57 mg N L−1) was recorded by Posadas et al. (2013) during RDW treatment in an algal-bacterial biofilm reactor at similar 5.2-day HRT and pH 7.0, while our results showed a maximum elimination of ≈ 13 mg N L−1. However, stripping was probably the main removal mechanism in the system of authors, while assimilation into biomass was the only mechanism of nitrogen removal in the present HRAP. Thus, the lower TN– RE is a consequence of the high nitrification activity combined with the lower pH, preventing N–NH4+ removal by stripping. As a result, the TN effluent concentrations unfortunately remained above the TN of 15 mg N L−1 established in Directive 98/15/CEE (European Union 1998) for wastewater discharge into the environment. Likewise, P effluent concentrations also remained higher than P concentrations permissible for wastewater discharge into the environment (2 mg P L−1). Influence of CO2 source on removal efficiency of pathogens and total coliforms The use of microbial indicators such as Pseudomonas aeruginosa, enterococci, Escherichia coli, and total coliforms now extends beyond indicating potential health risks due to fecal contamination (El-Khateeb et al. 2009). Bahlaoui et al. (1997) reported lower P. aeruginosa removal efficiency (91.6 ± 2.3%; ≈ 1.1 log–REs) in a 48-m2 HRAP treating raw wastewater, at 8day HRT, when compared with stage I of our experiments. Furthermore, the increasing of P. aeruginosa RE at stage II (≈ 99.6% and 2.5 log–REs) can be related with CO2. In this context, Eklund (1984) studied the effect of CO2 concentration (5– 80%) on the inhibition of the P. aeruginosa growth rate, obtaining a positive linear correlation (≈ 5–80% of REs). Furthermore, it is known that other environmental conditions like DO, sunlight intensity, and pH can affect the bacterial removal efficiency (Curtis et al. 1992; Heubeck et al. 2007). However, according to
Pearson’s correlation, the major factor responsible for the P. aeruginosa REs was the CO2 addition (Pearson’s r of 0.961, e.g., correlation of ≈ 96%), while DO, sunlight intensity, and pH had no statistical influence (p < 0.05). It must be stressed though that despite the statistical analysis showed no significant effect of sunlight intensity on P. aeruginosa REs at 95% of significance, higher Pearson’s r correlation of 0.699 (e.g., correlation of ≈ 70%) suggests that sunlight intensity may have also influenced the P. aeruginosa REs. In this case, more studies dealing on the P. aeruginosa removal and the influence of sunlight intensity are necessary, which is a suggestion for more investigations. Similar enterococci removal efficiencies were obtained during the HRAP operation, despite the CO2 at stage II. Lower enterococci REs (≈ 2.0 log–REs) were recorded by Awuah et al. (2002) in an algae-based domestic treatment pond at 7-day HRT and pH of ≈ 7.4. According to Ouali et al. (2013), the kinetic coefficient of E. coli removal only increases when the pH exceeds a value of 8.5, higher than our results (6.8–7.7). In this context, the lower E. coli REs obtained by Heubeck et al. (2007) and Posadas et al. (2015) can be related with higher temperatures or lower HRTs (Reinoso et al. 2008; Ouali et al. 2013). For instance, Posadas et al. (2015) recorded E. coli REs of ≈ 90% (1.0 log–REs) in an 850-L HRAP treating domestic wastewater at similar pH 7.0 (controlled by CO2 from flue gas), but at a temperature of 22 °C and 3-day HRT. Likewise, the maximum log–RE of ≈ 2.0 was found by Heubeck et al. (2007) in batch experiments with a 3day duration treating domestic wastewater in a HRAP with CO2 supply, at 21–22 °C. In this way, more studies dealing the mechanisms of E. coli removal are essential. The reductions obtained in our experiments were enough to achieve a final effluent that would meet the WHO guidelines on the fecal bacteria limit (≤ 1000 fecal coliforms per 100 mL) for the use of treated wastewater in unrestricted agricultural irrigation (WHO 2006). Reinoso et al. (2008) recorded similar results for total coliforms (≈ 2.4 log–REs) in a 1717-m3 facultative pond, but at 75.9-day HRT. According to Pearson’s correlation, the main factor responsible for the total coliform REs was the CO2 addition, rather than DO, sunlight intensity, or pH. Garcia-Gonzalez et al. (2007) reported that an aqueous medium with the addition of CO2 can inhibit microbial activity due to damage of the decarboxylation reactions in cells. On the other hand, some bacterial species are more tolerant to exposure to CO2 than others (Salih 2011), which can explain the fact that the P. aeruginosa and total coliform removal efficiencies were influenced by CO2 addition, while the effect on the enterococci and Escherichia coli removal efficiencies was negligible. Productivity and settleability of biomass, microalgae population, and elemental composition The similar light irradiances recorded at stages I and II contributed to obtaining comparable TSS concentrations and
J Appl Phycol
biomass productivities during both stages. Contextually, Posadas et al. (2015) obtained a higher productivity of ≈ 14 g m−2 day−1 in an outdoor HRAP treating RDW at a similar pH 7.0 controlled by CO2 from flue gas, but at much higher irradiance (462 ± 225 W m−2). Removal efficiencies of TSS at stage II were similar to the TSSSet–REs of 90 ± 15% found by Posadas et al. (2014a) in a comparable HRAP treating fishery and domestic wastewaters. Furthermore, Park and Craggs (2010) point out that CO2 addition in HRAPs can promote aggregation/bioflocculation of the microalgae with bacterial flocs to further increase biomass settling. Contextually, the increasing settleability at stage II can also be influenced by the predominance of filamentous Microspora sp., which improved the formation of flocs (Zhang et al. 2016). In this respect, the possibility of P. aeruginosa and total coliform removal by attachment to microalgal-bacterial flocs should be stressed, since these microorganisms have a tendency to adhere (Bitton 2005; ElKhateeb et al. 2009). On the other hand, the changing algal community from stages I to II can be associated with the decreasing pH (addition of CO2) and/or DO, increasing organic loading rate or even inhibition of C. vulgaris due to toxin production by Microspora sp. (Rodolfi et al. 2003; Pham et al. 2014; Sutherland et al. 2016). At this point, it can be stressed that these toxins may also inhibit P. aeruginosa. Thus, more studies focusing on the actual mechanisms of pathogen removal and changes in the algal community in HRAPs with a CO2 supply are necessary. Finally, the elemental composition obtained was in agreement with previous literature findings (C 40–60%; N 4–9%) (Grobbelaar 2003; Serejo et al. 2015). Conclusions The use of 30% CO2 at 5 mL min−1 in a 180-L HRAP for microalgae-based secondary domestic wastewater treatment resulted in an the Pseudomonas aeruginosa removal efficiency increasing from 97.4% (1.6 ± 0.1 log–REs) to 99.6% (2.5 ± 0.5 log–REs) and in total coliform removal efficiency increasing from 88.7% (1.1 ± 0.1 log–REs) to 99.4% (2.8 ± 1.2 log–REs). In contrast, the enterococci (≈ 99.7%; 2.6 log–REs) and E. coli (≈ 98.6%; 2.2 log–REs) removal efficiencies were not significantly altered, but were higher than those reported in the literature for the microalgalbacterial process. Acknowledgements The authors wish to thank CNPq - Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq, Brazilian National Research Council) for the financial support.
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